Carbon sources and sinks in high-elevation spruce–fir forests of the Southeastern US

Forest Ecology and Management 238 (2007) 249–260 www.elsevier.com/locate/foreco

Carbon sources and sinks in high-elevation spruce–fir forests of the Southeastern US H. Van Miegroet a,*, P.T. Moore a, C.E. Tewksbury a,1, N.S. Nicholas b a

Utah State University, Department of Wildland Resources, 5230 Old Main, Logan, UT 84322-5230, United States b Division of Resources Management and Science, Yosemite National Park, El Portal, CA 95318, United States Received 17 August 2006; received in revised form 18 October 2006; accepted 18 October 2006

Abstract This paper examines carbon (C) pools, fluxes, and net ecosystem balance for a high-elevation red spruce–Fraser fir forest [Picea rubens Sarg./ Abies fraseri (Pursh.) Poir.] in the Great Smoky Mountains National Park (GSMNP), based on measurements in fifty-four 20 m  20 m permanent plots located between 1525 and 1970 m elevation. Forest floor and mineral soil C was determined from destructive sampling of the O horizon and incremental soil cores (to a depth of 50 cm) in each plot. Overstory C pools and net C sequestration in live trees was estimated from periodic inventories between 1993 and 2003. The CO2 release from standing and downed wood was based on biomass and C concentration estimates and published decomposition constants by decay class and species. Soil respiration was measured in situ between 2002 and 2004 in a subset of eight plots along an elevation gradient. Litterfall was collected from a total of 16 plots over a 2–5-year period. The forest contained on average 403 Mg C ha1, almost half of which stored belowground. Live trees, predominantly spruce, represented a large but highly variable C pool (mean: 126 Mg C ha1, CV = 39%); while dead wood (61 Mg C ha1), mostly fir, accounted for as much as 15% of total ecosystem C. The 10-year mean C sequestration in living trees was 2700 kg C ha1 year1, but increased from 2180 kg C ha1 year1 in 1993–1998 to 3110 kg C ha1 year1 in 1998–2003, especially at higher elevations. Dead wood also increased during that period, releasing on average 1600 kg C ha1 year1. Estimated net soil C efflux ranged between 1000 and 1450 kg C ha1 year1, depending on the calculation of total belowground C allocation. Based on current flux estimates, this old-growth system was close to C neutral. # 2006 Elsevier B.V. All rights reserved. Keywords: Nutrient cycling; Management; Global climate change; Carbon budget; Coarse woody debris; Soil organic matter

1. Introduction There is considerable debate about the role of old-growth and unmanaged forests in the global carbon (C) cycle (Ciais et al., 1995). Currently, northern mid-latitude forests function as net C sinks for atmospheric carbon dioxide (CO2) (Pacala et al., 2001; Houghton, 2003). While commercial plantations in the eastern United States are net C sinks (Battle et al., 2000; Schimel et al., 2000), the role of mature or old-growth forests is less straightforward and more variable, and can range from C

* Corresponding author. Tel.: +1 435 797 3175; fax: +1 435 797 3796. E-mail address: [email protected] (H. Van Miegroet). 1 Present address: University of Arizona, School of Natural Resources, 325 Biological Sciences Building, P.O. Box 210043, Tucson, AZ 85721-0043, United States. 0378-1127/$ – see front matter # 2006 Elsevier B.V. All rights reserved. doi:10.1016/j.foreco.2006.10.020

neutral, to small net C sources or sinks (Pregitzer and Euskirchen, 2004). The C dynamics in old-growth forests differ from those in aggrading forests in several ways: (i) total C pools in live and dead wood are generally high in old-growth forests; (ii) net biomass increment declines in forest ecosystems that carry large above-ground woody biomass in large trees of older age classes; (iii) net ecosystem C exchange declines as net primary productivity tapers off (Harmon et al., 1990; Pregitzer and Euskirchen, 2004). Limited logging in some areas may further result in a natural build up of standing and downed wood in the ecosystem, which have the potential to release C via heterotrophic respiration (Harmon and Hua, 1991), and can ultimately determine the overall C source or sink strength of the system (Pregitzer and Euskirchen, 2004). Even in managed systems, age-related patterns in net ecosystem C fluxes are strongly driven by disturbance, reflecting the balance between inputs and decomposition of harvesting residues versus C

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increment in the young regrowth (Law et al., 2003; Pregitzer and Euskirchen, 2004). Large amounts of C also accumulate in the forest floor and upper mineral horizons of temperate and boreal forests, especially in absence of catastrophic fires (Pregitzer and Euskirchen, 2004). Small changes in this surficial soil organic carbon (SOC) pool due to climate variations can results in significant CO2 fluxes (Raich and Tufekcioglu, 2000; Rustad et al., 2000). Particularly the labile fraction of SOC appears to be temperature-sensitive (Fang et al., 2005), although some studies have shown the SOC turnover to be insensitive to temperature (Liski et al., 1999; Giardina and Ryan, 2000). A legacy of large detrital C and SOC pools and/or a recent disturbances, especially when coupled with limited growth potential, could potentially drive forest ecosystems to increase their source strength and become a net source of CO2. The cool temperate, high-elevation red spruce–Fraser fir [Picea rubens Sarg./Abies fraseri (Pursh.) Poir] forests in the southeastern US are an example of mature coniferous forests with large standing ecosystem C stocks. They are relic islands of a contiguous north–south chain of vegetation that existed during the Pleistocene (Delcourt and Delcourt, 1984) and currently occupy around 27,000 ha in southwestern Virginia, eastern Tennessee, and western North Carolina, with the largest area (74%) located in the Great Smoky Mountains National Park (GSMNP) (Dull et al., 1988). These spruce–fir ecosystems primarily consist of mature forest and 60% have not experienced severe, stand-replacing disturbances such as fire or logging (Harmon et al., 1983; Pyle and Schafale, 1988). The prevalence of old unglaciated landscapes with a legacy of soil C and limited fire history (Pyle, 1988) has resulted in C-rich forest soils (Johnson et al., 1991; Joslin et al., 1992). A combination of gap dynamics typical of mature forests and an infestation by the exotic balsam woolly adelgid (Adelges picea Ratz., BWA), causing extensive mortality of mature fir since the 1970s, has created a heterogeneous forest structure, with large variations in forest composition, stand age, live and dead standing biomass, and downed coarse woody debris (CWD) (Nicholas et al., 1992; Pauley et al., 1996; Rose, 2000). To date, limited data is available on C dynamics in these forests. It is our hypothesis that the combination of ecology and geology, historical and more recent disturbance regimes, and the current designation as a low-management area (national park or wilderness area) limits the ability of these forests to store additional C in the short-term. Furthermore, regional climate forecasts for the southeastern US include increases in temperature (Cooter, 1998) and a change in precipitation regime (Mearns et al., 2003). Accelerated turnover of existing detrital and SOC pools could cause these ecosystems to become a net source of CO2 under future regional climate change scenarios. The objectives of this analysis are (1) to determine the magnitude and distribution of C pools in spruce–fir forests typical of the high-elevation zone in the southeastern US; (2) to estimate major C fluxes into and out of the major ecosystems compartments; (3) to evaluate the current C balance of this spruce–fir ecosystem. There are several approaches to quantify the net ecosystem exchange of CO2, such as ground-based estimates of pool

changes, micrometeorological and eddy covariance measurements, and simulation modeling, each with their own spatial and temporal inference space (Curtis et al., 2002; Harmon et al., 2004). In ecosystems where the terrain is not flat, as is the case with most forests in mountainous areas, direct measurement of net ecosystem CO2 exchange with eddy covariance methods is impossible (Baldocchi, 2003), and more indirect methods have to be used. Long-term forest inventory data allow for relatively easy estimation of changes in live and dead wood C pools (e.g., Smith et al., 2004), but accurately estimating a change in the SOC pool is much more problematic. It requires a vast sampling effort in order to detect a temporal trend among data that inherently have large spatial variability (Wilding et al., 2001), and where changes may be small in comparison to pool sizes. In our analysis we use a variation of the ‘‘aggregated’’ flux method described by Harmon et al. (2004), combining changes in growth and mortality derived from periodic stand inventories, and estimates of annual heterotrophic respiration rates from the decay of standing and downed dead wood, with soil C fluxes estimated from shorter-term litterfall and in situ soil respiration measurements. In this paper we will use the term ‘‘C sequestration’’ specifically to designate the net rate of storage within a given compartment. 2. Materials and methods 2.1. Study site and plot layout The study was conducted in the spruce–fir zone of GSMNP in the proximity of Clingman’s Dome, on the North Carolina and Tennessee border. The core of our measurements were conducted in and adjacent to the Noland Divide Watershed (NDW) (358340 N lat., 838290 W long.), a small (17.4 ha), highelevation (1700–1900 m) first-order gauged catchment that typifies the southern Appalachian red spruce–Fraser fir forest type. It capitalized on existing infrastructure and available data from prior and ongoing research on forest dynamics (Nicholas et al., 1992), nutrient cycling (Johnson et al., 1991; Van Miegroet et al., 1993), and watershed processes (Nodvin et al., 1995; Van Miegroet et al., 2001). Mean annual air temperature is 8.5 8C, ranging from an average of 2 8C in January to +18 8C in July, with a frost-free period from May through September. Mean annual precipitation (MAP) is approximately 230 cm, ranging from 150 to 300 cm, and is fairly evenly distributed throughout the year (Van Miegroet et al., 2001). Snow accounts for about 10% of the MAP and typically covers the ground for 50 days a year (Johnson et al., 1991). The growing season is relatively short (100–150 days), and the climate in the spruce–fir zone is characterized by frequent cloud immersion and high winds (White and Cogbill, 1992). During the Integrated Forest Study (IFS), higher fog immersion was observed above 1800 m (Johnson and Lindberg, 1992), roughly matching the elevation of Clingman’s Dome road which bisects the watershed into an upper and lower section (Fig. 1). This is an uneven-aged forest, with red spruce and Fraser fir as the dominant overstory species. Species distribution follows

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Fig. 1. Location map of the Noland Divide Watershed and the four NAPAP plots in the Great Smoky Mountains National Park.

an elevational gradient. The overstory is dominated by mature, red spruce in the lower elevations (1370–1650 m) and the understory consists mainly of hardwoods including yellow birch (Betula lutea Michaux f.), mountain-ash (Sorbus americana Marshall), and mountain maple (Acer spicatum Lam.). At middle elevations (1675–1890 m), red spruce and Fraser fir codominate. At higher elevations (>1890 m), the forest has undergone canopy decline due to Fraser fir mortality caused by the infestation of the BWA, and the overstory is dominated by young and standing dead Fraser fir with a component of red spruce (Nicholas et al., 1992). The soils are Inceptisols with spodic characteristics classified as Dystrochrepts or Haplumbrepts, underlain by Thunderhead Sandstone (Johnson et al., 1991; Van Miegroet et al., 1993). They have a silt loam to sandy loam texture, and are generally shallow throughout the NDW (<50 cm depth to bedrock, Van Miegroet et al., 2001). During IFS, soil depth was estimated at 60–65 cm at nearby spruce–fir sites (Johnson and Lindberg, 1992). Soils are acidic, characterized by high organic matter content and low base saturation, as well as high nitrogen (N) mineralization and nitrification capacity (Johnson et al., 1991; Garten and Van Miegroet, 1994). In 1993, 50 permanent 20 m  20 m plots were established systematically along nine elevation bands ranging from approximately 1700 to 1900 m (Pauley et al., 1996) (Fig. 1). At the onset of the study, live basal area (LBA) of overstory trees [>5 cm diameter breast height (DBH)] in NDW ranged from 14.9 to 67.5 m2 ha1. Red spruce comprised 77% of LBA, while yellow birch and Fraser fir comprised 19% and 2.5%, respectively. Pauley et al. (1996) estimated average total biomass of live overstory trees for NDW at 220 Mg ha1. Fraser

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fir accounted for 70% of the standing dead stems, a pattern most likely related to the disturbance created by the BWA infestation, which caused the mortality of mature fir in the 1980s. The research site was later impacted by three hurricanes (Andrew in 1992, Opal in 1995, and Ivan in 2004) and by an ice storm in the winter of 1995 (Smith, 1997), which caused downing of live and dead trees and a significant input of CWD, especially in the upper part of the catchment (Van Miegroet et al., 2001). To encompass the full elevation range of the spruce–fir forest within the GSMNP (Nicholas et al., 1992), four additional 20 m  20 m plots were included in our design: two near Clingman’s Dome at an elevation above NDW (1966 m) and two below the elevation of the catchment outlet (1524 and 1536 m) (Fig. 1). They were selected from a network of spruce–fir forest inventory plots established as part of the National Acid Precipitation Assessment Program (NAPAP) (Nicholas et al., 1992) established in the mid-1980s. These four NAPAP plots and four plots within the NDW, two in the upper catchment (1835 m) and two in the lower catchment (1700 m), were designated as intensive plots for the study of soil C dynamics (Tewksbury, 2005; Tewksbury and Van Miegroet, 2006). Selection criteria for these eight intensive plots were: stratification along the full elevation range of the spruce–fir forest within the GSMNP (1525–1980 m); proximity to one another and similar aspect to minimize possible differences in geology and climate; relatively intact forest structure to minimize the confounding effect of recent gaps on soil microclimate and C dynamics; and overstory species composition and LBA representative of that particular elevation band (Nicholas et al., 1992). 2.2. Measurements and calculations 2.2.1. Carbon pools Overstory tree inventories were conducted at the NDW in 1993, 1998, and 2003, using protocols described by Zedaker and Nicholas (1990). In the four NAPAP plots, inventories were conducted in 1996 and 2002. The DBH, species, and crown position of each live and dead overstory (5 cm DBH) tree were recorded. Biomass of foliage, live branches, dead branches, bole bark, and bole wood of each tree was determined from DBH using allometric equations developed for red spruce, Fraser fir and yellow birch by Nicholas (1992). Biomass of other overstory species was estimated with the predictive equations of Weaver (1972). Carbon pools were calculated from biomass estimates and C concentration determined earlier from destructive sampling of biomass components for the major overstory species (spruce, fir and birch) within the NDW (Barker, 2000; Barker et al., 2002). The C content in live coarse roots was assumed to be 20% of aboveground tree C, based on estimates for cold temperate, needle leaf evergreen forests with Inceptisol-dominated soil orders (Vogt et al., 1996; Jackson et al., 1996). No belowground C estimates were made for dead snags. Shrub and tree stems with DBH >2 cm but <5 cm were surveyed in four 4 m  4 m subplots within each of the 54 permanent plots. Stem diameter was measured with a caliper

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15 cm above ground level in 2003; biomass of foliage and woody components was calculated using species-specific equations of Nicholas (1992) and Weaver (1972); and C content using C concentrations from Barker (2000). The stem diameter of seedlings <2 cm were measured in 16 1 m  1 m subplots within each of the 54 permanent plots with a caliper at 15 cm above ground level. Predictive biomass equations for individuals of this size were developed from destructive analysis outside the GSMNP. In the spruce–fir zone of the nearby Pisgah National Forest, 10 individuals of each species covering the observed stem diameter range were measured with a caliper, harvested, dried at 65 8C for 24 h, and weighed. Predictive equations for seedling biomass were derived from regressions between caliper size and dry weights by species using the PROC REG procedure in SAS (SAS, 2002). In the calculation of seedling C pools, an average tissue C concentration of 48% was assumed, which reflects the weighted average concentration earlier determined by Barker (2000) for the main tree species. Percent cover of all herbaceous plants (i.e., herbs, grasses, and mosses) was assessed by ocular estimation in 16 1 m  1 m subplots delineated with a PVC square frame inside each of the 54 permanent plots. To maintain the integrity of the permanent plots, species-specific regression equations relating percent cover to biomass were developed from destructive sampling in temporary 1 m  1 m plots established in Fall 2003 outside the permanent plots but still within the NDW. Inside these temporary plots, percent cover of each species was estimated to the nearest percent and specimens were clipped at ground level until adequate observations (n = 12) of all species of herbaceous plants (i.e., herbs, grasses, and mosses) were measured and collected. Plant samples were dried at 65 8C for 24 h, weighed, ground in a Wiley mill (# 40 mesh screen; 0.85 mm mesh), and subsamples were analyzed for C concentration using a LECO CN analyzer (CHN 1000, LECO Corp., St. Joseph, MI). Predictive equations for dry plant biomass were derived by regressing observed percent cover values with individual dry weights using the PROC REG procedure in SAS (SAS, 2002). These equations were used to calculate biomass of the herbaceous layer in each of the 54 permanent plots from cover herbaceous cover estimates. Because of the relatively short growing season (May through September) only one sampling period was necessary (Yarie, 1980), and the peak standing biomass at the end of the growing season was our estimate of aboveground herbaceous biomass. The C content of the standing dead trees (snags) was determined by species from the most recent inventory (2002 or 2003) using allometric equations of Nicholas (1992) and Weaver (1972) and C concentrations from Barker (2000). For trees first reported as standing dead in 1993, only the bole biomass was included, and current biomass was calculated assuming exponential decay of the snags using the decomposition constants of 0.006, 0.022, and 0.073 year1, respectively, for standing dead red spruce, Fraser fir and hardwood species reported in Rose (2000). For trees that were first reported as dead in the most recent forest inventory, bole, branch, and bark components were used in the calculation of C

content, and it was assumed that no significant C loss had occurred since death. This was done to differentiate trees that had only recently died and still had their branches and bark relatively intact from trees that had been standing dead for many years and where only the bole was likely to have persisted. The C pools in downed CWD were separated into old and recent inputs. The C content of old downed CWD was calculated for 2004, based on the initial biomass values in 1994 and a mean C concentration of 47% (Rose, 2000) and applying an exponential decay function for 10 years using decomposition constants by species and decay class reported for NDW by Rose (2000). Since only catchment-level CWD data were available for 1994, mean values were applied to all plots. The C content of new downed CWD was estimated for each individual plot from C content of the woody components of all live and dead standing trees that disappeared from the most recent forest inventory and were assumed to have fallen down. Carbon content of the O horizon (forest floor) in the 50 NDW plots was calculated from dry weight measurements from four replicate cores (inside diameter 4.1 cm) per plot taken in summer 1998 and 1999 as part of in situ N mineralization assays (Van Miegroet, unpublished data). In the eight intensive plots, O horizon material was removed in October 2003 from a 15 cm  15 cm area using a PVC sampling square laid out at four random locations within each plot. All samples were ovendried (65 8C), ground, and analyzed for C using a LECO CN analyzer. Mineral soil cores were taken at four locations at the periphery of each of the 54 plots (in summer 1999 for NDW; in summer 2004 for the eight intensive plots). Soils were sampled from 0 to 15 cm and then in 10 cm sections to a depth of 50 cm or until bedrock. Since the 0–15 cm core section represented a combination of O- and A-horizons, the organic horizon was removed and only the mineral soil was further analyzed. Samples were ovendried (105 8C), sieved (2 mm mesh) and crushed with a mortar and pestle, composited per depth increment and plot, and analyzed for total C on a LECO CN analyzer. Bulk density and percent gravel for mineral soil was determined using the core method (Blake and Hartge, 1986) in the eight intensive plots (four replicate cores per plot). Bulk density measurements from the eight intensive plots (four NAPAP, four NDW) were applied to the individual plots. Results from the two upper elevation intensive plots in the NDW were applied to the 17 other plots located above 1800 m in the catchment; those from the two lower elevation intensive plots were applied to the other 29 plots in the lower part of the catchment. The mineral soil C pools to a maximum depth of 50 cm were calculated from C concentration of the fine fraction, bulk density and gravel content of the corresponding soil sections. In the case where soils could not be sampled all the way to 50 cm, soil C content was calculated to the maximum depth of the last soil section that could be sampled. 2.2.2. Carbon fluxes To determine the net ecosystem C balance, the original formulation of Harmon et al. (2004) was expanded and

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separated into aboveground and belowground components, using the following formula (Law et al., 2003): NEP ¼ ðNPPAG  RhCWD Þ þ ðTBCA  RS Þ

(1)

where NEP is the overall net ecosystem C balance, NPPAG the aboveground overstory C sequestration, RhCWD the C release from decomposing standing and downed dead wood, TBCA the total belowground carbon allocation, and RS is the soil respiration. Throughout this paper, C fluxes from the atmosphere to the ecosystems (sinks) are given a negative sign; C fluxes from the ecosystem to the atmosphere (sources) a positive sign. NPPA: Net rate of C storage in the overstory was calculated retrospectively from net C increment in living trees and ingrowth between periodic tree inventories. Mean annual net C increment was defined as the increase in individual tree C content (i.e., individual tree growth) between inventory dates divided by the number of years between measurements. Trees that died or fell over between inventories were given a zero growth rate (as per Clark et al., 2001). Carbon sequestration in ingrowth was defined as the C content of all trees not previously inventoried as >5 cm. All C increment values were annualized based on a 10-year record (1993–2003) in NDW and on a 6-year data record (1996–2002) in the NAPAP plots. RhCWD: The annual C release from the standing and downed CWD was based on current C pool estimates described earlier and species-specific decay constants from Rose (2000). All trees classified as standing dead in the most recent inventory, were assigned species-specific decomposition constants for standing dead decay class 1 (0.006 year1 for spruce, 0.022 year1 for fir, 0.073 year1 for birch); new downed C input were assigned species-specific decay constants for downed CWD dead decay class 1 (0.036 year1 for spruce, 0.029 year1 for fir, 0.217 year1 for birch); and old downed CWD was assigned appropriate decay constant for downed wood by species and decay class (decay class 1: 0.036 year1 for spruce, 0.029 year1 for fir, 0.217 year1 for birch; decay class 2: 0.097 year1 for spruce, 0.067 year1 for fir, 0.084 year1 for birch; decay class 3: 0.270 year1 for spruce, 0.237 year1 for fir, 0.097 year1 for birch) determined earlier by Rose (2000) and also reported in Creed et al. (2004). All CWD categories were assumed to decay exponentially over a 10-year period in the future. Average annual heterotrophic respiration (C release) associated with decomposition of standing and downed CWD in each of the 54 plots was calculated by dividing the difference in initial and final C content by 10, and summing annualized C losses across species and CWD categories. RS: Soil respiration was measured in summer and fall in the eight intensive plots over a 2-year period for a total of 10 measurements between May 2002 and 2004. At four locations in each plot, CO2 efflux from forest floor and mineral soil was measured over a 24 h period using the static chamber technique with 2N NaOH as the trapping agent (Cropper et al., 1985; Raich et al., 1990), followed by backtitration with 0.75N HCl in the laboratory. Due to access problems, soil respiration measurements were not taken during the winter, and average daily winter respiration rates were estimated for each plot from

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measured summer rates and average soil temperatures using the equation (Zak et al., 1993): k1 ¼ k2 eðt1 t2 Þ=10 ln Q10

(2)

where k1 is the calculated mean winter respiration rate, k2 the average measured summer respiration rate, t1 the average winter soil temperature, t2 the average summer soil temperature, and Q10 = 2. Average soil temperatures were based on measurements at 1–2 h intervals using Stowaway Tidbit dataloggers (Onset Computer Corporation, Bourne, MA) placed at 10 cm soil depth in the center of each of the eight intensive plots. Data were organized into three periods; Summer: 15 May– 31 August; Fall: 1 September–15 November; Winter: 15 November–15 May. For each period, average daily respiration rate was multiplied by the number of days (Summer: 107 days; Fall: 76 days; Winter: 181 days) and annual CO2 emission from the soil was calculated as the sum of the seasonal values. Because passive trapping of CO2 with NaOH tends to underestimate actual soil respiration rates (Davidson et al., 2002; Knoepp and Vose, 2002), all mean annual CO2 fluxes were divided by an average methodological correction factor 0.465 obtained from Knoepp and Vose (2002) to approximate RS . TBCA: No direct root measurements were preformed in this study, and belowground C allocation was estimated indirectly for the eight intensive plots only, using four different approaches based on annual aboveground litterfall C and soil respiration fluxes. Litterfall was collected using round collection traps (0.07 m2) placed at the four corners of the four NAPAP plots and in 12 plots within the NDW (six each in the upper and lower part of the watershed). Samples were collected from the NAPAP plots in May, August, October for 2 years (2001–2003), and for the NDW during the same months over 5 years (1998–2003) as part of another study (Barker et al., 2002). All samples were oven-dried at 65 8C and weighed; composited by plot; sorted into four fractions: (1) needles, (2) foliage, (3) twigs and bark, and (4) other; ground in a Wiley mill (# 40 mesh screen; 0.85 mm mesh); and analyzed for C using a LECO CN analyzer. For the four NAPAP plots, plot-specific litterfall fluxes were used; the average litterfall across the six upper litterfall plots was applied to the two intensive plots in the upper NDW; the average litterfall across the six lower litterfall plots was applied to the two intensive plots in the lower NDW. The first approximation for TBCA (TBCA#1) consisted of multiplying average annual litterfall C flux by 1.85, intermediate between the conversion factors obtained by Raich and Nadelhoffer (1989) and Davidson et al. (2002). The second estimation for TBCA (TBCA#2) was simply based on the difference between RS and litterfall as per Raich and Nadelhoffer (1989) modified by Law et al. (2003), with the assumption that there were no significant changes in belowground C pools. The third method (TBCA#3) partitioned RS, based on the observation by Gaudinski et al. (2000) that O and A horizons generally account for two-thirds of total CO2 release from the soil. In our calculation we assumed that all CO2 efflux originating from below the O and A horizon and 30% of the

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CO2 originating from within the O and A horizon were rhizosphere derived. Finally, the fourth estimation (TBCA#4), applied the same calculation method but assumed that as much as 50% of the CO2 originating from within the O and A horizon were rhizosphere derived. These partitioning values between root and microbial CO2 efflux were based on general ranges cited on the literature for temperate and boreal conifer forests (Hanson et al., 2000; Bond-Lamberty et al., 2004; Vogel et al., 2005). Carbon balance: The aboveground component of the carbon balance was determined for each of the 54 plots, and averages and standard deviations were calculated across the lower NAPAP (n = 2), the lower NDW (n = 31), the upper NDW (n = 19) and the upper NAPAP plots (n = 2) (Table 2). The belowground part of the ecosystem C balance was determined

for the eight intensive plots only. Estimates of ecosystem C balance (NEP) were calculated from average below- and aboveground net C fluxes. 3. Results and discussion 3.1. Carbon pools The total average C content of the spruce–fir ecosystem in 2003 was 403 Mg ha1. More than half of that C was stored belowground in organic and mineral soil horizons and coarse root biomass (Table 1). The lowest ecosystem C capitals generally occurred in the upper NDW (<400 Mg C ha1), due to the combination of lower standing tree biomass and lower soil C. The latter was consistent with shallower soils to bedrock

Table 1 Carbon distribution in spruce–fir forests of the Great Smoky Mountains National Park (Mg ha1) in 2003 Lower NAPAP [1524–1537a; 2b]

Lower NDW [1701–1801a; 31b]

Upper NDW [1835–1911a; 19b]

Upper NAPAP [1951–1966a; 2b]

Ecosystem average [1524–1966a; 54b]

Overstory vegetation Abies—Live Abies—standing dead Betula—live Betula—standing dead Picea—live Picea—standing dead Other species—live Other species—standing dead

0.41  0.20 0.52  0.74 23.6  16.8 0 118  15.0 0.016  0.023 4.98  0.17 1.18  1.32

2.44  2.93 4.43  3.80 41.8  46.1 2.41  5.02 91.4  44.3 20.3  22.9 1.38  1.83 0.62  1.47

12.4  11.4 7.66  4.27 10.3  23.0 0.37  0.95 82.9  52.7 19.2  23.7 0.77  1.22 3.67  6.47

25.8  6.34 11.5  5.16 0 0 106  11.0 14.5  20.5 1.72  2.43 1.89  0.15

6.74  9.34 5.68  4.41 28.5  40.5 1.51  3.96 90  46.0 19.0  22.3 1.31  1.76 1.76  4.19

Total live Total standing dead Total overstory

147  31.7 1.72  0.56 149  31.2

137  49.2 27.8  22.5 165  49.9

106  48.9 30.9  24.6 137  48.8

133  2.18 27.9  15.5 164  17.7

126  49.3 27.9  23.2 155  49.2

Understory vegetation Woody (shrubs and conifers) Herbaceous

6.35  1.58 0.16  0.20

3.66  2.84 1.03  0.69

3.79  3.10 1.34  0.88

2.20  0.02 0.38  0.04

3.75  2.87 1.08  0.78

Total understory

6.51  1.38

4.68  2.75

5.13  3.46

2.58  0.02

4.83  2.96

Downed coarse woody debris Old downed woodc Abies Betula Picea

4.09 0.29 3.92

4.09 0.29 3.92

4.09 0.29 3.92

4.09 0.29 3.92

4.09 0.29 3.92

New downed woodd Abies Betula Picea

0.67  0.95 17.3  24.5 18.0  6.02

2.53  3.20 1.59  3.45 17.8  27.5

4.55  5.08 0.41  1.37 24.2  28.3

3.77  2.71 0.98  1.39 0

3.22  3.99 1.73  5.34 19.4.  26.9

Total coarse woody debris

44.2  17.5

30.2  27.7

37.5  28.7

13  1.3

32.7  27.5

Belowground roots

29.5  6.34

27.43  9.85

21.3  9.77

26.7  0.44

25.3  6.0

Forest floor (O)

30.5  7.66

21.6  6.78

18.2  5.13

22.5  8.79

20.8  6.6

Mineral soil (depth in cm) A (to 15 cm) A + B (15–50 cm)

24.4  8.14 166  20.4

39.7  11.5 130  43.7

35.2  11.7 115  36.1

41.3  16.2 164  11.5

37.6  14.5 127.4  58.3

221  19.86

191  46.2

168  40.0

228  18.9

185.7  50.5

451

419

370

432

403

Location

Total soil (0–50 cm) Ecosystem a b c d

Elevation (m). Number of plots Based on 1994 watershed-level survey by Rose (2000). Based on plot-level overstory inventories.

H. Van Miegroet et al. / Forest Ecology and Management 238 (2007) 249–260

at the upper slope positions due to frequent landslides (White and Cogbill, 1992). The soils in the NDW were on average <50 cm to bedrock (Van Miegroet et al., 2001), but 70% of the soils in the upper NDW had shallower depth to bedrock compared to 45% of the soils in the lower catchment. The size and distribution of C pools and distribution were similar to values for this forest reported earlier in the IFS (Johnson and Lindberg, 1992). Carbon stored in aboveground live biomass (160 Mg ha1) was less than half the value reported by Harmon et al. (2004) for old-growth forests in the western Cascades. Forest floor C pools (21 Mg ha1) were at the low end of estimates for other montane conifer forests in North America (30–60 Mg ha1, Simmons et al., 1996; Quimet et al., 1996; Tremblay et al., 2002; Smith and Heath, 2002), while mineral soil C pools (190 Mg ha1) tended to be in the higher range of reported values (60–180 Mg ha1, Edmonds and Chappell, 1994; Tremblay et al., 2002; Pregitzer and Euskirchen, 2004). This may be due to the fact that it was often difficult to discern purely organic forest floor from organic-rich mineral soils, especially in ecosystems like the spruce–fir, where mixing of layers due to windthrow and landslides has occurred (Federer, 1982). Our soil C pool sizes were higher than the mean calculated by Miller et al. (2004) for well-drained soils in southwestern Virginia (112 Mg ha1), but closely corresponded to the average of 201 Mg C ha1 estimated by Kern (1994) for Haplumbrepts, the dominant soil classification in the spruce–fir. They were similar to the worldwide average for wet boreal forests (Post et al., 1982; Callesen et al., 2003).

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Across the entire dataset the total C in standing and downed CWD was similar, 28 and 33 Mg ha1, respectively. Cumulatively, CWD represented almost 15% of the entire ecosystem C. Fir represented only a small fraction (5%) of the live overstory C pool, but was on average 20% of the C in standing dead trees, with a high of 40% in the highest elevation NAPAP plots (Table 1). Downed CWD increased since the 1994 (18 Mg ha1, Rose, 2000) and the 1998 inventories (22 Mg ha1, Creed et al., 2004) indicating the effect of more recent forest disturbances. Understory constituted only a small C pool (<5 Mg ha1). All C pools were spatially variable, with variability lowest for forest floor and mineral soil (CV = 30%) and highest for the dead and live overstory pools (CV = 79% and 39%, respectively) (Table 1), reflecting the spatial heterogeneity in canopy structure caused by repeated disturbances over the past three decades. The increase in live and dead fir C with increasing elevation and the concomitant decline in birch was typical of shifts in forest composition with elevation (Nicholas et al., 1992; Pauley et al., 1996). 3.2. Carbon fluxes Despite heavy mortality of mature fir and recent windfalls, C was still accumulating in living trees, albeit at a low rate (mean 2700  820 kg C ha1 year1), the majority by spruce (Table 2). Aboveground C sequestration rates varied in space (CV = 30% across 54 plots) with the contribution by fir increasing and that of birch declining with elevation (Table 2).

Table 2 Carbon fluxes in spruce–fir forests of the great smoky Mountains National Park (kg C ha1 year1) Location

Lower NAPAP

Aboveground C sinks Overstory NPP Abies Betula Picea Ingrowth Total overstory sequestration

16  6 486  8 2457  127 49  44 3033  169

Upper NDW

Upper NAPAP

93  138 632  653 1703  672 209  167 2637  672

643  659 155  231 1312  862 480  338 2590  917

1588  1095 0 2401  1393 511  412 4499  114

+167 +20 +207

+167 +20 +207

+167 +20 +207

+167 +20 +207

C release new downed wood Abies Betula Picea

+17  24 +1533  2167 +543  182

+64  81 +141  306 +539  831

+115  128 +36  121 +733  856

+95  68 +87  123 0

C release from standing dead Abies Betula Picea

+13  18 0 0

+109  93 +196  370 +121  135

+185  104 +29  88 +117  138

+280  125 0 +85  120

C sources C release old downed wood Abies Betula Picea

Decomposition coarse woody debris

Lower NDW

+2500  1944

+1564  808

+1609  868

+941  49

533  1775 +1794  221 +1261

1074  1172 +1639  185 +565

981  1401 +1373  309 +392

3559  163 +1001  331 S2557

Net aboveground C fluxes Soil C fluxes Ecosystem C balance + sign indicates C source and  sign indicates C sink.

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Fig. 2. Average annual C sequestration in tree increment and ingrowth for the various inventory periods [significant differences using repeated measures ANOVA in total C sequestration, increment, and ingrowth between inventory periods for upper and lower NDW is indicated by ***p < 0.001 and **p < 0.01].

Since the early 1990s, C sequestration rates in aboveground woody biomass significantly increased (Fig. 2; Moore et al., 2006) from an overall mean of 2180 kg C ha1 year1 in the period 1993–1998 to an average of 3110 kg C ha1 year1 in the period 1998–2003. This increase in C storage rates was particularly pronounced at higher elevations, where disturbance effects were typically heaviest (Pauley et al., 1996). Ingrowth represented an additional C sink of 310 kg C ha1 year1, 74% of which (229 kg C ha1 year1) associated with fir seedlings. The role of ingrowth in C sequestration generally increased with elevation (Fig. 2), with a tenfold difference between the lower and upper NAPAP plots (Table 2, Fig. 2), reflecting the greater impact of stand-opening disturbances at higher elevations. Total aboveground C sequestration rates for this forest (2700 kg C ha1 year1) were similar to those reported for old-growth coniferous forests in similar climates (Gordon, 1981; Harmon et al., 2004; Pregitzer and Euskirchen, 2004). The upswing in growth of mature trees over the last 5 years, the increasing contribution of fir in wood increment (from 9% between 1993 and 1998 to 16% between 1998 and 2003), and the considerable ingrowth dominated by fir seedling recruitment, all signal that the forest in general and fir in particular were recovering from the BWA infestation disturbance. Our findings did not support the notion of a growth decline in this spruce–fir forest predicted earlier by McLaughlin et al. (1987) and McLaughlin and Tjoelker (1992). The role of understory as a C sink was small and unlikely to increase considerably. Assuming annual turnover of all herbaceous and about half of the woody understory C pool and accumulation of the remainder of the woody understory C over a 10-year period (Table 1), the net C increment was <200 kg C ha1 year1. Understory was therefore not further included in our ecosystem C balance calculations. On the other hand, we may have overestimated C sequestration rates due to ingrowth as they were based on seedling C pools at entry in the inventory divided by time between inventories, instead of by the unknown period of time it took stems to reach the 5 cm DBH threshold. Given these opposing trends, we assumed that current C storage rates calculated for ingrowth represented a reasonable estimate of annual C sequestration rates in all

woody species with stems <5 cm (i.e., spruce and fir regeneration and understory shrubs). Decomposition of the sizable detrital (standing and downed) wood C pool represented a net C source of 1600  640 kg C ha1 year1, mostly from logs that fell in the last 10 years. Plotlevel data were highly variable, consistent with the patchiness of disturbance, but averages for the upper and lower NDW were surprisingly similar. The average C release from old downed wood and standing dead trees was 400 kg C ha1 year1. Dead fir accounted for most of the C release from standing CWD, and their contribution increased with elevation (Table 2). There are few literature values against which to compare our CWD C release rates. Harmon et al. (2004) reported a C release of 660 kg C ha1 year1 from dead snags, 670 kg C ha1 year1 from downed logs, and 530 kg C ha1 year1 from fine woody debris, or a total of 1860 kg C ha1 year1 in mixed conifer forests in Western Oregon, close to the values reported here. The C release rates reported by Marra and Edmonds (1994, 1996) for decomposition of downed Douglas-fir and western hemlock logs in a temperate, old-growth rain forest in Olympic National Park were higher: in the order of 4400 kg C ha1 year1 for decay class 1–2; 3700 kg C ha1 year1 for decay class 3; and 4200 kg C ha1 year1 for decay class 5. Chambers et al. (2001) estimated a C release rate from coarse wood decay of 1900 kg C ha1 year1 in a central Amazon old-growth forests, while Law et al. (2003) reported significantly lower C fluxes (250–550 kg C ha1 year1) from decomposing wood in mature and old ponderosa pine stands in central Oregon. Annual estimated soil respiration (RS) rates ranged from 2600 to 3900 C ha1 year1 across the eight intensive sites, with an overall mean of 3140 kg C ha1 year1. Our RS values were lower than the values for some common temperate coniferous forests reported by Kane et al. (2003) (5000–7000 kg C ha1 year1), Rustad et al. (2001) (3000–12,000 kg C ha1 year1), and Davidson et al. (2002) (6500–9900 kg C ha1 year1). They were higher than soil respiration rates for fir in New Brunswick (1130 C ha1 year1) and red pine in New York (1500 kg C ha1 year1) (Raich and Tufekcioglu, 2000), and comparable to CO2 efflux rates for spruce in Alaska summarized by Raich and Tufekcioglu (2000)(3100 kg C ha1 year1) and Vogel et al. (2005) (3100–6000 kg C ha1 year1). One potential reason for the difference between our and published soil respiration rates is that winter CO2 efflux rates were estimated from summer rates and temperatures, assuming a Q10 = 2, rather than being measured directly. Also, the actual Q10 factor can vary significantly across a wide range of climatic conditions (Raich and Schlesinger, 1992; Peterjohn et al., 1994), ranging from 1.6 to 3.5 (Kirschbaum, 1995; Boone et al., 1998), with the Q10 values generally higher at lower temperatures (Davidson and Janssens, 2006). Finally, static chambers with passing trapping agents tend to underestimate CO2 evaluation rates (Davidson et al., 2002) and correction factors must be empirically derived (Knoepp and Vose, 2002). Since no other RS values are available for this forest type our values constitute the best available estimates. Average annual litterfall C fluxes gradually decreased from 1800 kg C ha1 year1 in the lower NAPAP plots, to 1600 and

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Fig. 3. Mean annual soil respiration (RS) and belowground C allocation (TBCA) averaged across the eight intensive sites. For explanation of different calculation methods of TCBA see text.

1400 kg C ha1 year1 in the lower and upper NDW, to a low of 1000 kg C ha1 year1 in the highest NAPAP plots. Averaged across collection points and years, mean annual litterfall C return for this spruce–fir forest was 1450 kg C ha1 year1, similar to averages measured in this forest in the late 1980s (Johnson et al., 1991). The first TBCA estimate, based on a 1:1.85 ratio between aboveground litterfall and TBCA (Raich and Nadelhoffer, 1989; Davidson et al., 2002) was considerably higher than the three others (Fig. 3). It likely overestimated TBCA, especially in view of the shallow soils (<50 cm) (Van Miegroet et al., 2001), and the use of a global ratio (see Davidson et al., 2002). The convergence in the three values obtained from site-specific data was encouraging, and suggested some robustness in our estimates. The average annual TBCA based on the difference between RS and aboveground litterfall (TBCA#2), was identical to the mean value (1700 kg ha1 year1) based on spatial (66% of RS from O and A; 34% from deeper mineral soil) and functional separation (all deep respiration and 30% of shallow respiration is rhizosphere-derived) of soil respiration (TBCA#3). Assuming all deep and half of the shallow portion of soil respiration was rhizosphere-derived (TBCA#4), resulted in a slightly higher value of 2100 kg ha1 year1 (Fig. 3). 3.3. Ecosystem carbon balance Based on the difference between aboveground NPP and C release from woody detritus decomposition, the average aboveground C sequestration rate of this spruce–fir forest was 1100 kg C ha1 year1. The results were spatially highly variable (CV = 110%), and plot-level values ranged between a net C release of 720 and a net C sequestration of 3700 kg C ha1 year1. There was, however, no distinguishable pattern with elevation. The upper NAPAP plots stood out by their high net C sequestration rates (3600 kg C ha1 year1), the result of high net growth (Fig. 2) and low CWD accumulation rates (Table 1), and suggesting that these plots remained relatively unaffected by spruce blowdowns that were especially prevalent near the top of the ridge.

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Given the similar outcomes of TCBA#2 and TBCA#3 (Fig. 3), these fluxes seemed most appropriate for deriving net C release rates from the soil, which averaged 1450 kg C ha1 year1 (Table 2). This was similar the average (1200 kg C ha1 year1) across mature and old ponderosa pine stands in a Central Oregon chronosequence (Law et al., 2003). The belowground contribution to the ecosystem C balance was subject to considerably more uncertainty than the aboveground estimates, given the lack of direct root measurements, and the need to estimate TBCA. Also, fluxes used to estimate net soil C release were generally measured at a different temporal and spatial resolution than aboveground fluxes. Assuming belowground steady state conditions, net soil C efflux was primarily associated with decomposition of fresh litterfall, which accounted for 45% of RS, compared to the 30% reported in Raich and Nadelhoffer (1989). As a whole, the ecosystem was close to C neutral, with a mean net sequestration of 500 kg C ha1 year1 across the eight intensive plots and a net release of 400 kg C ha1 year1 for the entire 54 plots. This was consistent with the metadata analysis of Pregitzer and Euskirchen (2004), which showed boreal forests and older temperate forest to hover around zero NEP, and with the small sink sizes of old-growth Douglas-fir (430 kg C ha1 year1) and ponderosa pine forests (350 kg C ha1 year1) reported by Harmon et al. (2004) and Law et al. (2003), respectively. As was the case with the contributing C fluxes, the ecosystem C balance varied spatially, even among plots within the same elevation band, and appeared to change gradually with elevation. Spatial variability in the NEP of old-growth forests has been attributed mostly to changes drivers of soil- and CWDderived heterotrophic respiration (Pregitzer and Euskirchen, 2004). Law et al. (2003) argued that the soil was the main source of heterotrophic respiration, as the decomposition of CWD was relatively slow at their dry sites. In our system, where decomposition processes were not limited by lack of moisture (Tewksbury, 2005), and large amounts of CWD had accumulated in the wake of the BWA infestation, CWD decomposition was a significant contributor to the net ecosystem C balance (Table 2). The lower-elevation sites emerged as stronger net C sources, while highest elevation sites acted as net C sinks, due greater post-disturbance C sequestration in wood increment and ingrowth combined with lower heterotrophic respiration rates in colder soils (Tewksbury, 2005; Tewksbury and Van Miegroet, 2006) (Table 2, Fig. 3). The greatest degree of uncertainty in our C balance was associated with the estimation soil respiration and TBCA. Zerva et al. (2005) argued that belowground steady state conditions, which underly the estimation of TBCA from soil respiration and aboveground litterfall (Raich and Nadelhoffer, 1989) may not be met in disturbed forests. However, the extensive soil coring required to obtain more direct measurements of fine root dynamics and the destructive nature of such sampling was prohibitive at our field sites (National Park). While we were reasonably confident in our litterfall estimates, temporally and spatially more extensive respiration data (using different measurement techniques) might have resulted in a

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more refined assessment of the ecosystem C balance. Considering that our average total heterotrophic respiration rate (3000 kg C ha1 year1), from decomposition of aboveground litter (1450 kg C ha1 year1) and CWD (1590 kg C ha1 year1) was much lower than the 5770 kg C ha1 year1 reported old-growth forests in Oregon (Harmon et al., 2004), it is quite possible that this high-elevation forest already functioned as a net source of C. The future trajectory of the C balance for these ecosystems remains uncertain. Heterotrophic C sources could increase under future climate change scenarios. Indeed, several field studies (Trumbore et al., 1996; Trumbore, 1997; Kane et al., 2003; Emmett et al., 2004) including at our intensive study plots (Tewksbury, 2005; Tewksbury and Van Miegroet, 2006) have shown and increase in soil respiration with increasing soil temperature unless soils get too dry (Emmett et al., 2004). Garten et al. (1999) also showed more rapid turnover of SOC with increasing temperature across an elevation gradient in southern Appalachian forests. Also, the forest dynamics in this system are typically driven by the catastrophic disturbance associated with the mortality of mature fir by BWA and subsequent windthrow of spruce (Moore et al., 2006), and the decomposition of this CWD is likely to remain an important and potentially increasing source of CO2 in the future. On the other hand, opening of the forest canopy has already caused increased growth of remaining trees and accelerated release of saplings in the understory (Moore et al., 2006). It is unclear to what extent current post-disturbance growth rates can be maintained in the future as continuing waves of BWA infestation and mortality are expected. 4. Conclusions Many of the results in this analysis of C pools and fluxes were consistent with our traditional understanding of structure and function of old-growth forest systems. Our study also pointed out that in evaluating the role of oldgrowth ecosystem in the global C balance, a clear distinction needs to be made between C sequestration in terms of total amount of C accumulated in the past (which is typically large) and C sequestration rates, as the net balance between autotrophic C assimilation rates and heterotrophic C releases to the atmosphere. In forests that contain large amounts of legacy C as SOC and CWD, turnover of these pools may critically affect the ecosystem C balance, which, under climate change scenarios, may shift towards a greater net source of CO2. Considerable resources are currently being expended into regional assessment of forest ecosystem C pools (e.g., Smith et al., 2004), and C balances. Yet few studies have quantified the role CWD dynamics with empirical field data. Ecosystem C balance calculations in mature forests with limited management should include this component because it may constitute a significant source of CO2 possibly equivalent to or exceeding soil respiration (Marra and Edmonds, 1994, 1996; Chambers et al., 2001; Table 2). Little is currently known about CWD decomposition rates under changing environmental conditions.

Furthermore, there is still uncertainty about changes in forest disturbance regimes caused by global change, and the extent to which post-disturbance growth might be able to compensate for the CO2 release from the decay of old and newly deposited CWD. Based on our estimates it is difficult to conclude with a great degree of confidence whether this spruce–fir system is still C neutral, or has already become a net C source. In our study area, the lower elevation sites appeared to be stronger C sources. Forest regrowth following recent disturbances increased C sink strength at the highest elevations, where mortality was highest and heterotrophic respiration was more temperature-limited. The future trajectory may be quite different as it is uncertain how the long current post-disturbance growth pulse can be maintained and to what extent soil respiration will increase in the future. Acknowledgements We graciously thank Alan Mays, Larry Shelton, Anita Rose and Suzanne Fisher from the Tennessee Valley Authority (TVA) for all of their logistical assistance; Michael Griffiths, Bryan Drew, Faye Tewksbury, and Lisa Moore for their help in the field; numerous students in the USU Forest Soils Lab for their help in preparing and processing field samples; and Louise O’Deen and Banning Starr from the US Forest Service Rocky Mountain Research Station for the chemical analyses. We also thank TVA and EPA for the use of the forest inventory data sets, and Mike Ebinger for his assistance with the development of the GIS maps. Funding for data collection was provided by the USDA National Research Initiative Competitive Grants Program (Grant No. 97-35101-4314 to Utah State University), the USGS Biological Research Division (Cooperative Agreement No. 1434 HQ97-RV-01555 RWO27 and RWO34 to the Utah Cooperative Fish and Wildlife Research Unit), and Tennessee Valley Authority’s Public Power Institute. AES Publication No.7676, Utah Agricultural Experiment Station, Utah State University, Logan, Utah 84322-4810. References Baldocchi, D.D., 2003. Assessing eddy covariance technique for evaluating carbon dioxide exchange rates of ecosystems: past, present and future. Global Change Biol. 9, 479–492. Barker, M., 2000. The Role of Overstory Nitrogen Uptake as a Sink of Atmospheric Deposition in Southern Appalachian Spruce–Fir Forests. MS Thesis. Utah State University, Logan, Utah. Barker, M., Van Miegroet, H., Nicholas, N.S., Creed, I.F., 2002. Variation in overstory nitrogen uptake in a small, high-elevation southern Appalachian spruce–fir watershed. Can. J. Forest Res. 32, 1741–1752. Battle, M., Bender, M.L., Tans, P.P., White, J.W.C., Ellis, J.T., Conway, T., Francey, R.J., 2000. Global carbon sinks and their variability inferred from atmospheric O2 and d13. Science 287, 2467–2470. Blake, G.R., Hartge, K.H., 1986. Bulk density. In: Knute, A. (Ed.), Methods of Soil Analysis. Part 1. Physical and Mineralogical Methods. Agronomy Series 9. 2nd ed. American Society of Agronomy, Madison, pp. 363–382. Bond-Lamberty, B., Wang, C., Gower, S.T., 2004. A global relationship between heterotrophic and autostrophic components of soil respiration? Global Change Biol. 10, 1756–1766.

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